Aquatic photochemistry of the sulfonamide antibiotic sulfapyridine

Journal of Photochemistry and Photobiology A: Chemistry 262 (2013) 14–21

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Aquatic photochemistry of the sulfonamide antibiotic sulfapyridine Jonathan K. Challis a,b , Jules C. Carlson a,b,c , Ken J. Friesen b , Mark L. Hanson c , Charles S. Wong a,b,∗ a

Department of Environmental Studies and Sciences, Richardson College for the Environment, University of Winnipeg, Winnipeg, Manitoba R3B 2E9, Canada Department of Chemistry, Richardson College for the Environment, University of Winnipeg, Winnipeg, Manitoba R3B 2E9, Canada c Department of Environment and Geography, University of Manitoba, Winnipeg, Manitoba R3T 2N2, Canada b

a r t i c l e

i n f o

Article history: Received 22 February 2013 Received in revised form 12 April 2013 Accepted 14 April 2013 Available online 22 April 2013 Keywords: Direct and indirect photolysis Sulfonamides Sulfapyridine Quantum yield Photodegradation products

a b s t r a c t The photolytic behavior of the sulfonamide antibiotic sulfapyridine in water was investigated using a laboratory photoreactor approximating full-spectrum sunlight. Direct photolysis of sulfapyridine was rapid, with a half-life of 2.6 h and 31 min, and an observed quantum yield of 0.0013 ± 0.0002 and 0.013 ± 0.001, for the neutral species and fully deprotonated species, respectively. Direct photolysis rates increased dramatically with degree of deprotonation, with measured pKa1 and pKa2 values of 2.22 ± 0.03 and 8.58 ± 0.02, respectively. Indirect photolysis was assessed using water from constructed wetland mesocosms. A four-fold increase in the dissipation rate of sulfapyridine was observed due to the influence of high levels of dissolved organic carbon, after accounting for light screening by such materials. Nitrates had no observable effect on indirect photolysis rates. Major photoproducts identified were SO2 extrusion and OH addition products. These results show that photolytic processes are a major removal mechanism of sulfonamide drugs in aquatic systems. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Human and veterinary-use pharmaceuticals are a significant class of contaminants that can enter aquatic environments. These chemicals find their way into surface waters primarily via wastewater discharge or agricultural runoff, and are normally not completely removed in sewage treatment processes. Although acute effects from such chemicals are generally unlikely at the low concentrations prevalent in the environment, little is known of the potential effects of sublethal chronic exposure concentrations on non-target organisms, particularly antibiotics [1]. These drugs are among the most prominent therapeutic pharmaceuticals detected in the environment [1], and are of particular concern due to the potential for development of microbes bearing antibiotic resistance genes [2–5]. Sulfonamides are a major class of broad-spectrum synthetic antibiotics, widely used to treat respiratory, skin, and urinary tract infections, and as animal growth promoters. This class of chemicals are excreted in urine and have been detected at concentrations as high as 7.9 ␮g/L in raw wastewater [6], with sulfapyridine observed in effluent at 63–135 ng/L [7]. The strong bacteriostatic properties of sulfonamides can have significant effects on the ecological

∗ Corresponding author at: Richardson College for the Environment, University of Winnipeg, Winnipeg, Manitoba R3B 2E9 Canada. Tel.: +1 204 786 9335; fax: +1 204 774 2401. E-mail addresses: [email protected], [email protected] (C.S. Wong). 1010-6030/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jphotochem.2013.04.009

functioning of microorganisms. For example, sulfapyridine was shown to significantly reduce microbial activity in soils, with EC10 values ranging from 0.14 to 160 ng/g based on the Fe(III) reduction test, with an incubation time of 7 d [8]. This, along with the promotion of antibiotic resistance [2,3], strongly suggest that the environmental presence of sulfonamides may pose a hazard to ecosystem health. Sulfonamides absorb light in the environmentally relevant UVB and UV-A ranges (280–400 nm) [9]. Thus, direct photolysis could be a significant mechanism for abiotic transformation in sunlit surface waters, as observed for other five- and six-member ringed sulfonamides for which protonation state has a major influence on degradation rate [9,10]. Furthermore, the presence of photosensitizing species can significantly enhance the extent of indirect photolysis. Dissolved organic matter (DOM) present in natural waters can sensitize chemical photodegradation either by direct transfer of energy, or through the formation of reactive intermediates such as singlet oxygen (1 O2 ), superoxide anion (O2 −• ) and hydroxyl (• OH), hydroperoxy (HO2 • ), alkylperoxy (RO2 • ), and carbonate (CO3 −• ) radicals [9,11–15]. Additionally, dissolved nitrates can mediate the formation of • OH radicals in natural waters, and in some cases to a greater degree than DOM [16]. While a number of studies have examined the photochemical fate of sulfonamides, few have investigated sulfapyridine, despite its common occurrence [7]. Photodegradation rates of sulfapyridine differed in wastewater treatment plant effluent compared to in pure HPLC-grade water [17]. Degradation rates were 5-fold faster in the effluent (t1/2 = 2 h) versus in the pure water in that study (t1/2 = 10 h), indicating that indirect photolysis plays a significant

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role in limiting the persistence of sulfapyridine. Sulfamethazine, on the other hand, showed little evidence of indirect photolysis in the effluent matrix [17]. Quantum yields were not reported for either chemical. Despite the structural similarities that exist within the class of sulfonamide drugs, photochemical behavior differs significantly from compound to compound [9,10], suggesting that the study of the photochemical fate of individual sulfonamides on a case-by-case basis is necessary. The objective of the present study was to characterize the photochemical behavior of the sulfonamide antibiotic drug sulfapyridine. Specifically, direct photochemical degradation rates and quantum yields of sulfapyridine were examined in water with a laboratory photoreactor. The effect of speciation on direct photolysis was examined by conducting irradiations at acidic, neutral and basic pH. Indirect photolysis was assessed by irradiating sulfapyridine in water from field mesocosms designed to replicate constructed wetlands, that can help eliminate such chemicals [18]. Two photodegradation products were identified. These results will help define the fate of sulfapyridine in the aquatic environment. Furthermore, this study will further our understanding of the removal mechanisms responsible for eliminating pharmaceuticals from natural and engineered waters, and characterize exposure routes, therefore enhancing risk assessment of this chemical in aquatic ecosystems.

3.5 ␮m). Mobile phases were water and ACN buffered with 0.05% formic acid (10 mM) to pH 3. Sulfapyridine was separated isocratically with 70:30 H2 O:ACN, at 1 mL/min, while PNA/PYR was separated with 80:20 H2 O:ACN at 0.5 mL/min. Quantitation was monitored at 260 and 320 nm for sulfapyridine and 320 nm for PNA/PYR. Identification of photoproducts was done via ultra-high performance liquid chromatography–tandem mass spectrometry (UHPLC/MS/MS) using an Agilent 1200 binary UHPLC pump coupled to a 6410 triple quadrupole MS/MS (Agilent Technologies). Separation was achieved with a Zorbax C18 column (2.1 mm × 100 mm, 1.8 ␮m) with a Phenomenex SecurityGuard C18 Guard Cartridge (4 mm × 3.0 mm ID) using gradient elution at 0.5 mL/min commencing with 95:5 H2 O:MeOH (buffered with 10 mM formic acid), ramping linearly to 100% MeOH over 5 min, followed by a 3 min re-equilibration. Injection volume was 5 ␮L. Analytes were ionized using an electrospray interface operating in both positive and negative ionization modes with the following conditions: capillary voltage, 4000 V; nebulizer pressure, 15–55 psi; drying gas flow, 10–11 L/min; drying gas temperature, 300 ◦ C. The source fragmentor voltage ranged from −51 V in negative mode to +129 V in positive mode. Nitrogen was the collision gas, with collision energies between 11 and 20 V. The cell accelerator voltage was maintained at 7 V for all analyses.

2. Materials and methods

2.3. pKa determination

2.1. Chemicals and reagents

Spectra of sulfapyridine were recorded at 0.1–0.2 pH intervals between pH 12 and 1. Aqueous solutions were buffered with K3 PO4 (50 mM) titrated to the desired pH with concentrated HCl. At each pH a 10 mL aliquot of the buffer was removed and spiked with 100 ␮L of sulfapyridine providing a 5 mg/L solution. Differences in absorbance at single wavelengths were plotted as a function of pH and the resultant data were fit to a non-linear sigmoidal regression (Prism v. 5.01, GraphPad Software, La Jolla, CA) to determine pKa . Additional details are available in Supplemental Information.

Sulfapyridine (≥99%), p-nitroanisole (PNA, 97%), and pyridine (PYR, ≥99.9%) were purchased from Sigma–Aldrich (St. Louis, MO). HPLC grade methanol was purchased from Fisher Scientific (Fair Lawn, NJ). Solutions for direct photolysis were prepared with nanopure water (>17 M-cm, Milli-Q RG, Millipore Corp., Ann Arbor, MI), and buffered using di- and tri-basic potassium phosphate (K2 HPO4 , K3 PO4 , ≥98%, Sigma–Aldrich) titrated with hydrochloric acid (HCl, Fisher Scientific) to the desired pH. Ultra-grade potassium nitrate (KNO3 , ≥99.5%, Sigma–Aldrich) was used for indirect photolysis experiments. Liquid chromatographic solvents were prepared with nanopure water and HPLC grade acetonitrile (ACN, Fisher Scientific) or methanol (MeOH) buffered with formic acid (95%, Sigma–Aldrich). 2.2. Instrumentation and equipment 2.2.1. Laboratory photoreactor All irradiations were performed using a Rayonet Merry-GoRound Photochemical Reactor (model RPR-100, The Southern New England Ultraviolet Company, Branford, CT). The photoreactor had 16 medium-pressure mercury lamps with spectral emission ranging from 250 to 400 nm, centered at 300 nm. Irradiation vessels used were 50 mL cylindrical Pyrex tubes which filtered wavelengths <290 nm. 2.2.2. Chemical analyses A Shimadzu (Columbia, MD) UV-2501PC spectrophotometer was used to record UV–vis absorption spectra in the determination of molar absorptivities of sulfapyridine and pKa values via spectrophotometric titration. Analyses of irradiated solutions were done with an Agilent Technologies (Mississauga, ON) 1200 high performance liquid chromatograph (HPLC) equipped with a UV diode array detector. Sulfapyridine was separated on a Waters (Milford, MA) Symmetry C18 , 4.6 mm × 150 mm, 3.5 ␮m analytical column with a Phenomenex (Torrance, CA) SecurityGuard C18 Guard Cartridge (4 mm × 3.0 mm ID). PNA/PYR were separated using an Agilent Zorbax HILIC plus column (2.1 mm × 100 mm,

2.4. Photolytic studies 2.4.1. Chemical actinometry The PNA/PYR actinometer established by Dulin and Mill [19] was used to monitor photon flux, which was then used to estimate the photoreaction quantum yield of sulfapyridine. Photoreactor experiments used 4.6 × 10−5 M PNA and 0.01 M PYR. Irradiation and dark experiments for the actinometer system were carried out in parallel. 2.4.2. Direct photolysis Stock solutions of sulfapyridine were prepared at 4.0 × 10−4 M (100 mg/L) in nanopure water. Triplicate laboratory irradiations were conducted with 40 mL solutions of sulfapyridine, each at 5 mg/L, in Pyrex tubes. Dark experiments were carried out in an oven that matched the temperature and time of irradiation in the photoreactor, which reached a maximum temperature of 45 ◦ C at irradiation times >2 h. All experiments were performed in triplicate over five time points. Photolyses were conducted in 50 mM of the appropriate potassium phosphate buffer at pH 5.2, 7.2 and 11.1. 2.4.3. Indirect photolysis Laboratory photolysis of sulfapyridine was also investigated in simulated natural water from established field mesocosms. Absorbance spectra of the mesocosm water was obtained using a UV–vis spectrophotometer, and used to determine screening factors of the water. These simulated constructed wetland mesocosms contained sediment, water (originally sourced from City of Winnipeg tapwater), macrophytes, invertebrates, and microbial

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communities [20]. Mesocosm water from control tanks in which no sulfapyridine was spiked was filtered through 0.45 ␮m filters (Pall Corporation, Ann Arbor, MI) and adjusted to pH 7.2 with 20 mM phosphate buffer in order to prepare 5 mg/L sulfapyridine solutions for irradiations. The dissolved organic carbon (DOC) in mesocosm water was determined to be 15.7 ± 0.1 mg/L (n = 3) by ALS Laboratories (Winnipeg, MB). The addition of sulfapyridine and buffer diluted DOC concentrations to 15.2 mg/L in irradiation solutions. Nitrates were below detection limits (<0.01 mg/L). Carbonate alkalinity (CaCO3 ) was 248 ± 10 mg/L (n = 3) as determined using a standard test kit (LaMotte Company model DR-A, Chestertown, MD) [20]. Laboratory photolyses were conducted in triplicate for sulfapyridine in both unmodified mesocosm water, and that spiked with 3 mg/L KNO3 to test for the effect of DOC and NO3 − respectively. Nitrate concentrations were that of a typical small river system impacted by wastewater effluent inputs [21]. Samples were irradiated over sufficient time, across 12 time points to characterize degradation kinetics. Dark experiments, in an oven, were done in triplicate for mesocosm water and nitrate-spiked mesocosm water to account for any biotic and abiotic processes which could affect chemical losses. Photolyses were performed both with (1% v/v) and without MeOH present in solution in order to investigate the role of • OH in the indirect photolysis of sulfapyridine.

2.4.4. Photoproduct identification Solutions of 5 mg/L sulfapyridine in nanopure water adjusted to pH 7.2 with 20 mM phosphate buffer were irradiated in the photoreactor over two half-lives. Aliquots (3–5 mL) were purified by solid phase extraction using OASIS HLB (60 mg, 3 cc, Waters), eluted with 3 mL MeOH, evaporated to dryness with nitrogen gas, and reconstituted in a 90:10 H2 O:MeOH mixture to match the starting

eluent composition. This process separated background noise and potassium adducts from early-eluting peaks, and removed the phosphate buffer that would interfere with electrospray ionization. In some cases, depending on the concentrations of parent and product compounds, samples were diluted prior to analysis. Analysis was performed by UHPLC/MS/MS in full product scan mode. Molecular ions that appeared in only the mass spectra of the irradiated samples and not in that of the time-zero, dark, or blank samples were targeted for further investigation by optimizing the MS conditions for the specific masses.

3. Results and discussion 3.1. Effect of speciation on sulfapyridine light absorption Sulfapyridine absorbs light into the solar region ( > 290 nm) (Fig. 1), with a major absorption band at max = 309 nm (n → ␲*) in addition to a tailing absorption from the ␲ → ␲* transition with max = 261 nm. The amount of methanol (<1%, v/v) used to solubilize the compounds for UV–vis measurements in the determination of pKa should have an insignificant effect on the optical properties of sulfapyridine, since the refractive index of methanol is similar to that of water [14]. Aqueous speciation of sulfonamides is controlled by the pH and the acid–base characteristics of the compound’s amine ( NH2 ) and amide ( NH ) moieties (Fig. 1). Absorption spectra for sulfapyridine showed significant changes as the pH was varied between 1 and 12 (Fig. 1A and B). The observed pKa1 and pKa2 values, obtained from the inflection points in a nonlinear fit of the data, were 2.22 ± 0.03 and 8.58 ± 0.02, respectively (Fig. 1C and D), and in good agreement with the literature values determined experimentally; pKa1 = 2.6 and pKa2 = 8.4 [22].

Fig. 1. Speciation and absorption spectra of 5 mg/L sulfapyridine at selected pH values (A and B). Absorbance values (···) at 260 nm (for pKa1 ) and at 320 nm (for pKa2 ) plotted as a function of pH (C and D) and fitting of the data via non-linear sigmoidal regression (—). Reported uncertainty in pKa represents the 95% confidence interval of the inflection point. The absorption spectrum at pH 5 overlaps nearly completely with that at pH 7.

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The acidity of sulfapyridine’s amine group, quantified by pKa1 , falls within a relatively narrow range of 1.5–2.9, typical of sulfa drugs containing five- and six-member ring substituents. The pKa2 values of a series of sulfa drugs with different substituents varies from 4.7 to 8 [9,22]. More specifically, six-member rings with two N atoms that also contain CH3 groups have the greatest pKa2 values (ranging from 7 to 8), indicating that the electron-donating CH3 group somewhat offsets the effect of electron withdrawal by the N atoms, thus increasing the stability of the neutral amide and weakening its proton-donating ability. The second pKa of sulfapyridine is more basic than normally observed for sulfa drugs due to the presence of only one N atom in the heterocyclic substituent, which lessens electron withdrawal from the neutral amide (NH), decreasing its proton-donating ability. The pKa2 value is particularly relevant since it falls within the pH range of surface waters, which would therefore control the speciation of sulfonamides, dictate their light absorption characteristics and, potentially, affect their photochemical behavior. 3.2. Laboratory photoreaction studies 3.2.1. Photoreactor The elevated temperature of the photoreactor above 25 ◦ C (Tmax = 45 ◦ C at irradiations >2 h; Tave ≈ 35 ◦ C) is expected to have a small but insignificant effect on the acid–base equilibrium (pKa ) of sulfapyridine. Cross and Cao [22] studied the temperature effects on the acid–base speciation of a number of sulfonamides. While the data for sulfapyridine was not provided in the paper, they showed that the degree of dissociation varied from 1 to 4% for sulfadiazine and sulfadimethoxine, respectively, going from 30 to 40 ◦ C. Even assuming a ‘worst-case’ change of 10% in the overall speciation of sulfapyridine due to elevated temperatures, the effect would be insignificant since the irradiation solutions were buffered to pH values ≥2.5 pH units away from the pKa1 (2.22) and pKa2 (8.58) of sulfapyridine. Thus, the HS and S species will still predominate (≥98% assuming a 10% change) at pH 5.2 and 11.1 respectively, due to elevated temperatures in the photoreactor. 3.2.2. Chemical actinometry The quantum yield of the PNA/PYR actinometer was adjusted with 0.01 M pyridine (a = 0.437[PYR] + 0.000282) [19] to provide a half-life of approximately 15 min to match that of sulfapyridine in the photoreactor. The absorption spectrum of PNA overlapped the source emission to approximately 360 nm, thus, the measured pseudo-first-order direct photodegradation rate constant of the actinometer (kpa = 0.00025 s−1 ) is largely attributed to light absorption in the 290–360 nm range. Assuming the quantum yield (a = 0.0046) to be constant over this wavelength interval [19], the overall rate of light absorption is estimated from kp = a ·ka to be ka = 0.0543 s−1 . The wavelength-specific light absorption rates (ka,␭ ) may be estimated from wavelength-specific relative lamp emission and molar absorptivities of PNA (ka,␭ = I␭ εa,␭ ). Calculated ka,␭ values were then used in the estimation of source intensities (I␭ ) [14]: I␭ =

ka,␭ 2.303a εa,␭ l

(1)

Summing over the wavelength range (290–360 nm) provides the incident light intensity of (9.8 ± 0.5) × 10−7 E L−1 s−1 . Using an average estimated path length [23] of the Pyrex tubes (l = 2.78 cm), this intensity is equivalent to a light flux of (1.6 ± 0.2) × 1015 photons cm−2 s−1 . 3.2.3. Direct photolysis While small losses of 5–10% (maximum) were observed in dark experiments for sulfapyridine, these were not statistically

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Fig. 2. First-order direct photolysis of sulfapyridine buffered at three pH values for irradiations in a laboratory photoreactor and for the dark control. Error bars represent standard deviations of the mean.

significant given relative standard deviations of 9% in the slopes of the pseudo first-order decay curves (Fig. 2). Hence, dissipation rates were not corrected for other abiotic elimination. Sulfapyridine photodegradation was clearly pH-dependent (Fig. 2), indicating that speciation strongly affected sulfapyridine photolysis, as with other sulfonamides [9]. Sulfapyridine exists primarily as the fully deprotonated S species (fraction present, fs > 0.997) at pH = 11.1, while the singly-protonated species (HS) is most prevalent (fHS > 0.999) at pH = 5.2. The photodegradation rates of the target chemical (kpc ), sulfapyridine, observed at the above specified pH values are thus attributed to single species (Table 1). A change in pH from 11.1 to 5.2 resulted in a 5.8-fold decrease in photo-reactivity as speciation shifted toward the neutral, less reactive form of sulfapyridine. The direct photolysis rate constants for individual species (HS and S) allows for estimation of overall photodegradation rates (kp,est ) at pH values at which both species are present: kp,est = fHS · kHS + fS · kS

(2)

For sulfapyridine at pH 7.2 (fHS = 0.960 and fS = 0.040) the neutral but less reactive species (HS) is dominant, leading to an estimated photodegradation rate constant, kp,est = 0.0052 min−1 which is in reasonable agreement with the observed value of kpc = 0.0072 min−1 (Table 1). Photodegradation of sulfapyridine was slower at lower pH, despite the more favorable overlap of its absorption spectrum with the source emission spectrum. This is attributed to the lower quantum yield of the dominant HS species. Since the actinometer and sulfapyridine are irradiated in identical reaction vessels and under identical conditions, the reaction quantum yield for photodegradation of the chemical (c ) [14] was calculated (Table 1):



c =

kpc kpa



˙ I εa,␭ ˙ I␭ εc,␭



a

(3)

where c and a refer to the chemical and actinometer, respectively; kp is the observed photodegradation rate constant; ε␭ , and I␭ are the molar absorptivity and source intensity at wavelength ; and a is the actinometer reaction quantum yield. The value for S was 0.013 ± 0.002, based on pH 11.1 data. A tailing absorption was observed for S from 290 to 360 nm (Fig. 1), suggesting that a single electronic transition was responsible for light absorption and hence photodegradation. Therefore, S− should be constant over the wavelength region utilized [23]. At pH 5.2, the neutral form (HS) was dominant (fHS = 0.999), with a calculated reaction quantum yield HS of 0.0013 ± 0.0001. The reaction quantum yield (Table 1) for the neutral species was significantly less than for the fully deprotonated species (S ≈ 10HS ), which more than compensates

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Table 1 Direct photolysis reaction rate constants (kp ), correlation coefficients (r2 ), half-lives (t1/2 ) and quantum yields (c ) determined for the species HS, S, and at pH 7.2 of sulfapyridine in buffered solutions, using the laboratory photoreactor. pH

Species

kp (min−1 )a

r2

t1/2 (min)

c b

5.2 7.2 11.1

HS HS ↔ S S

0.0045 ± 0.0003 0.0072 ± 0.0001 0.022 ± 0.001

0.983 0.999 0.999

154 96 31

0.0013 ± 0.0001 N/A 0.013 ± 0.001

a b

Errors represent 95% confidence intervals. Errors calculated through error propagation. N/A – not applicable since multiple species were present.

for the greater absorptivity of HS to decrease photodegradation rates with decreasing pH. To our knowledge, these are the first experimentally observed reaction quantum yields reported for sulfapyridine. We do not expect indirect photolysis to play a significant role in these direct photolysis experiments. Indirect photolysis experiments (Section 3.2.4) showed • OH to have no apparent effect on the removal of sulfapyridine. Singlet oxygen, while present in the incubations from dissolution of atmospheric oxygen, would likely be at concentrations well below the 10−15 to 10−12 M in natural waters [24], given the lack of DOM or other photosensitizing species to mediate their formation. Given typical bimolecular second-order rate constants for reaction of 1 O2 with sulfonamides of 104 to 108 M−1 s−1 [9], the absolute maximum reaction rate that is even plausible (10−4 s−1 ) would be very small compared to that of direct photolysis (Table 1). Additionally, carbonate radicals may be present from partitioning of atmospheric CO2 (steady state [CO2 ] ≈ 1.3 × 10−5 M). However, concentrations are expected to be orders of magnitude below those found in surface waters ([CO3 −• ] ≈ 10−15 to 10−14 M) [25]. With bimolecular reaction rates in the range of 106 to 109 M−1 s−1 [26], contributions from CO3 −• should also be insignificant in water without significant amounts of precursors for indirect photolytic species. Our observed relation between molar absorptivities of different acid–base species of sulfapyridine and changes in quantum yield has also been observed for other sulfonamides; of which vary considerably even between very similar sulfonamides [9]. For example, sulfamoxole and sulfisoxazole differ only in the position of the nitrogen in the ring substituent (oxazole – N3 and isoxazole – N2, respectively), yet they exhibited opposite trends with regards to the relationship between molar absorptivity and quantum yield. Despite these structural similarities amongst the sulfonamide compounds, the significant differences in photochemical behavior observed among them suggest that the governing parameter affecting the photochemistry is the differing heterocyclic functional groups. Given the broad spectrum of possible functional groups, it is difficult to compare our results for sulfapyridine to other sulfonamides. Garcia-Galan et al. [17] is the only other study to our knowledge that investigated the photochemical behavior of sulfapyridine. They observed a half life of 11 h for sulfapyridine in HPLC grade water, determined using a SunTest CPS instrument equipped with a xenon arc lamp. However, they did not report quantum yields and since different light sources were

used, differences between rate constants or half-lives are difficult to interpret. Our observations for sulfamethoxazole (data not shown) are in general agreement with the literature [9,27,28] suggesting the methods used in this study are consistent with previous work. Specifically, our observations, as with other studies, showed that the neutral form (HS) of sulfamethoxazole is more photoreactive than the fully deprotonated form (S) with HS ≈ 6S− for this sulfonamide (data not shown). However, direct comparisons are confounded by differences in experimental design since light sources, source emission spectra, solution pH and the actinometer system used may all affect reported quantum yields [29]. 3.2.4. Indirect photolysis No significant loss of sulfapyridine was observed in any dark experiments. Hence, observed degradation rates were affected only by photolytic processes. Photolysis of sulfapyridine was rapid in mesocosm water (Table 2) but showed no significant change in the presence of 3 mg/L (≈48 ␮M) NO3 − suggesting that nitratemediated hydroxyl radical (• OH) production was not responsible for observed losses at these nitrate concentrations. Although it has been shown that nitrate-mediated • OH pathways only become important at nitrate concentrations above 100 ␮M (6.2 mg/L) [30], the nitrate concentrations used in this study were representative of a rural river system impacted by the release of sewage lagoons [21]. Furthermore, the presence of methanol, a known • OH quencher [31,32], had no measurable effect on observed photodegradation rates, further supporting the absence of an • OH mediated degradation by nitrate. Since DOC was a major constituent of mesocosm water, its role on the observed photolytic behavior of sulfapyridine was assessed by comparing photodegradation rates in mesocosm water (ktotal ) with those in nanopure water. As mesocosm water had a favorable absorption spectrum in the 290–360 nm region (Supplemental data, Fig. S2), light attenuation due to DOC was expected. A wavelength-specific screening factor (S␭ ) was estimated [33]: S␭ =

1 − 10−˛␭ l 2.303 ˛␭ l

(4)

where ˛␭ (cm−1 ) is the wavelength-specific attenuation coefficient (considered as the absorbance of DOC for the selected wavelength when the path length is 1 cm) and l (cm) is the path length of the irradiated Pyrex tubes. Mesocosm water had a relatively large

Table 2 Pseudo first-order photoreactor photolysis rate constants, half-lives (t1/2 ), and correlation coefficients (r2 ) for sulfapyridine in mesocosm water (ktotal , total photolysis; measured) and in buffered nanopure water – corrected for light screening (kp,corr , direct photolysis; calculated) – from laboratory photoreactor experiments at pH 7.2. Nanopure water

Mesocosm water

kp,corr (min−1 )a , b

ktotal (min−1 )a

r2

t1/2 (min)

kindirect (min−1 )c

% IPd

0.0028 ± 0.0001

0.012 ± 0.001

0.998

57

0.009 ± 0.001

75

a b c d

Errors represent the 95% confidence intervals. kp,corr = S␭ × kp,est . Indirect photolysis rate constant (kindirect ) calculated from ktotal = kp,corr + kindirect . % Indirect photolysis (IP): kindirect /ktotal.

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light screening factor ranging from 0.45 to 0.69 in the 290–360 nm region, similar to what Lam and Mabury [34] observed. An averaged light screening factor, S␭ , was determined by summing S␭ over the wavelengths 290–360 nm and normalizing to the absorbance of sulfapyridine in that region. S␭ = 0.55, meaning the average photon fluence rate was 55% of near-surface rates, suggesting that, in the absence of other effects, photodegradation rates should be decreased by the presence of DOC in natural water due to light screening. This effect has been observed for sulfamethoxazole [34,35] and confirmed in our laboratory (ktotal /kpc ≈ 0.86, data not shown). Application of S␭ to direct photolysis reaction rates determined in lab water provide an upper-limit estimate of expected direct photolysis rates in natural waters under identical irradiation conditions. This correction for light attenuation can be done using the equation: kp,corr = S␭ × kp,est

(5)

where kp,corr and kp,est are the corrected (accounting for light screening) and estimated (calculated using Eq. (2)) direct photolysis rate constants respectively, and S␭ is the average light screening over the wavelength region 290–360 nm. Using kp,est instead of kp,obs in this case allows for a more accurate estimate when working in the sensitive pH range between pKa1 and pKa2 of sulfapyridine (i.e. pH 7.2). For example, a change of 0.2 pH units above or below 7.2 will lead to a variation of 5–7% in kp,est . Thus, by using Eq. (2), the kp,est can be calculated at the exact pH that was used for the indirect photolysis experiments. The presence of 15.2 mg/L DOC resulted in an increased overall removal of sulfapyridine, with indirect photolysis accounting for 75% of the total photolysis (Table 2; ktotal /kp,corr ≈ 4). kp,corr ranged from 0.0023 to 0.0035 at the two extremes of the screening factor (i.e., at 290 and 360 nm respectively) with an intermediate value of 0.0028 (Table 2), calculated with S␭ . Garcia-Galan et al. [17] observed similar results when comparing the irradiation of sulfapyridine in HPLC lab water versus waste water effluent. They reported a fivefold increase in pseudo first-order photolysis rate constants for sulfapyridine between HPLC water and wastewater effluent agreeing well with results observed in the present study; and suggesting a significant contribution to photodegradation due to indirect photolysis mechanisms. The specific indirect photolysis mechanisms responsible for the increased removal rate of sulfapyridine in mesocosm water are not completely clear in the present study (Table 2; kindirect = 0.009 ± 0.001 min−1 ). While an • OH mediated mechanism is not supported, it is conceivable that other reactive species such as singlet oxygen (1 O2 ) generated through photo-processes involving excited states of DOM, or triplet excited state DOM (3 DOM) itself, may be responsible for enhanced photodegradation of sulfapyridine. Boreen et al. [10] attributed the enhanced degradation in natural waters of similar sulfa-drugs containing six-member ring heterocyclic groups to interaction with 3 DOM, although they also noted that 1 O2 may be contributing a maximum of 5 to 17% to the enhanced degradation of the sulfonamides studied. The source of the DOM, autochthonous (aquatic) or allochthonous (terrestrial), can have significant effects on the photo-mechanisms responsible. Guerard et al. [33] were able to show that aquatic DOM will primarily mediate degradation through 3 DOM intermediates, while terrestrial DOM is more reactive in promoting degradation by reactive oxygen species. Moreover, further studies have shown that DOM is capable of inhibiting the excited triplet-induced oxidation of many aromatic aquatic contaminants, specifically those containing aniline functionalities (e.g. sulfonamides) [36,37]. This inhibition has been attributed to anti-oxidant moieties (i.e. phenolic groups) present in DOM; with allochthonous DOM expected to more efficiently inhibit triplet-induced oxidation over less aromatic

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DOM (autochthonous) [36]. Our mesocosms were seeded with aquatic and riparian plants [20], suggesting that autochthonous DOM may play a role in inhibiting triplet-induced oxidation. That having been said, the sediments of our mesocosms were originally clean commercial potting soil [20] providing DOM likely of allochthonous origin. Further research is needed to delineate indirect photolysis mechanisms and pathways of contaminants in shallow surface-water wetlands. Due to the competing affects that DOM can have on photolysis in a system; photosensitization, light screening, scavenging, and oxidative inhibition, identifying the specific indirect photolysis mechanism(s) is non-trivial. Further research is required to understand fully the mechanism(s) responsible for the indirect photolysis of sulfapyridine. Nonetheless, as in nanopure water, the observed photolysis rate of sulfapyridine in mesocosm water is rapid. This observation suggests that natural humic substances likely contributed significantly to the overall photodegradation of sulfapyridine, at least under our specific incubation conditions. 3.3. Natural sunlight estimations Persistence of sulfapyridine under natural sunlight was estimated using L␭ values for 50◦ N latitude during mid-summer and fall [14] using the following equation: sun kp,est = c



εc,␭ L␭

(6)

␭

Estimated direct photolysis half-lives of sulfapyridine ranged from 4 h in summer to 12 h in fall. Although these estimations provide useful data in assessing the potential environmental fate of these compounds, factors such as the use of tubes rather than flat surfaces, the potential for light attenuation by water depth and natural chromophores dissolved in the water, and temperature gradients in the water column would also need to be considered in developing an accurate description of true environmental photochemical behavior. Nonetheless, these estimations agree closely with experiments conducted in shallow-depth outdoor mesocosms simulating constructed wetlands, where a total degradation half life of 5.5 h was observed for sulfapyridine during summer months [20]. This study did not identify specific degradation mechanisms, however based on the physical and chemical properties of sulfapyridine, and the results from the present study, photolysis is expected to dominate the removal of sulfapyridine from the water column. While the water in our incubations (Table 2) came from these outdoor mesocosms, other factors likely contributed to the lack of observed indirect photolysis in the latter, e.g., greater potential screening by DOM given the much deeper water column in the mesocosms compared to our incubation experiment. 3.4. Photoproduct identification Identifying photoproducts is necessary to completely understand the fate and effects of a chemical, as they have been shown to significantly contribute to a chemical’s toxicity [38]. Two major photoproducts were identified through our MS analysis (Supplemental data, Table S1) for sulfapyridine irradiation: an SO2 extrusion corresponding to m/z 186.1 [M + H − 64]+ and an OH addition product corresponding to m/z 266.1 [M + H + 16]+ . Both of these masses correspond to two of the major photoproducts previously identified for sulfapyridine [17]. Furthermore, SO2 extrusion is consistent with previous findings [10] that identified such products as primary photoproducts for a number of other six-member heterocyclic ring-containing sulfonamides. Further evidence confirming the identity of this photoproduct is based on the fact that extrusion of SO2 is a common photochemical process occurring in cyclic

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J.K. Challis et al. / Journal of Photochemistry and Photobiology A: Chemistry 262 (2013) 14–21

molecules [39]. Although monitoring and quantification of the formation of photoproducts over time was beyond the scope of the present study, it was evident from the data that as the irradiation time increased the abundance of the m/z 186.1 photoproduct increased (data not shown), further identifying it as a true photoproduct. Additionally, the ratio between signals of the product ion (m/z 108) and qualifier ion (m/z 169) of the SO2 -extrusion product (m/z 186.1) was approximately 4:1 across all irradiated samples; with the m/z 108 being four times as sensitive as the m/z 169. The sulfapyridine photoproduct corresponding to the m/z 266.1 is likely an OH-addition product. Evidence from our photolytic studies suggest that • OH does not play a role in the degradation of sulfapyridine, thus the mass addition of 16 [M + H + 16]+ suggests two possible mechanisms, not involving • OH. Addition of OH− to the double bond, followed by a rearrangement to reform the double bond with loss of a proton produces the m/z 266.1 ion; similar to the mechanism proposed for sulfamethoxazole [40]. Hydroxyl addition to the primary or secondary amine is also possible [17]. 4. Conclusions The present study investigated the relevance of direct and indirect photochemical processes on the persistence of the sulfonamide antibiotic drug sulfapyridine in aquatic systems. Sulfapyridine experienced rapid direct photolysis varying significantly with pH. At high pH, sulfapyridine, predominantly in its S form, exhibited enhanced photoreactivity compared to its HS form. Although the absorptivity of sulfapyridine increased as solutions became more acidic, the reaction quantum yield decreased as speciation shifted toward the neutral form. The present study reported the first experimentally measured quantum yields of sulfapyridine. Investigations in natural water illustrated that indirect photolysis mechanisms are likely responsible for a significant portion of the photochemical degradation of sulfapyridine in the environment. While • OH mediated degradation was ruled out through quenching experiments, the resulting increase in removal of sulfapyridine may be due to interaction with both 3 DOM and 1 O2 . Two major photodegradation products of sulfapyridine were identified in the present study; an SO2 extrusion and a hydroxylated product. Overall, this investigation provides new understanding of the photochemical behavior of sulfonamide drugs in aquatic environments. Acknowledgements We thank the Canada Research Chairs Program, the Natural Sciences and Engineering Research Council of Canada, the Lake Winnipeg Basin Stewardship Fund of Environment Canada, and the Thomas Sill Foundation for funding, and Jennifer Low for helpful discussions. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jphotochem. 2013.04.009. References [1] L.H.M.L.M. Santos, A.N. Araújo, A. Fachini, A. Pena, C. Delerue-Matos, M.C.B.S.M. Montenegro, Ecotoxicological aspects related to the presence of pharmaceuticals in the aquatic environment, Journal of Hazardous Materials 175 (2010) 45–95. [2] M.A. Gilliver, M. Bennett, M. Begon, S.M. Hazel, C.A. Hart, Enterobacteria—antibiotic resistance found in wild rodents, Nature 401 (1999) 233–234.

[3] A. Pruden, R. Pei, H. Storteboom, K.H. Carlson, Antibiotic resistance genes as emerging contaminants: studies in northern Colorado, Environmental Science and Technology 40 (2006) 7445–7450. [4] W. Witte, Medical consequences of antibiotic use in agriculture, Science 279 (1998) 996–997. [5] W. Zhang, B.S.M. Sturm, C.W. Knapp, D.W. Graham, Accumulation of tetracycline resistance genes in aquatic biofilms due to periodic waste loadings from swine lagoons, Environmental Science and Technology 43 (2009) 7643–7650. [6] X. Peng, Z. Wang, W. Kuang, J. Tan, K. Li, A preliminary study on the occurrence and behavior of sulfonamides, ofloxacin and chloramphenicol antimicrobials in wastewaters of two sewage treatment plants in Guangzhou, China, Science of the Total Environment 371 (2006) 314–322. [7] A. Göbel, C.S. McArdell, M.J.-F. Suter, W. Giger, Trace determination of macrolide and sulfonamide antimicrobials, a human sulfonamide metabolite, and trimethoprim in wastewater using liquid chromatography coupled to electrospray tandem mass spectrometry, Analytical Chemistry 76 (2004) 4756–4764. [8] S. Thiele-Bruhn, I.-C. Beck, Effects of sulfonamide and tetracycline antibiotics on soil microbial activity and microbial biomass, Chemosphere 59 (2005) 457–465. [9] A.L. Boreen, W.A. Arnold, K. McNeill, Photochemical fate of sulfa drugs in the aquatic environment: sulfa drugs containing five-membered heterocyclic groups, Environmental Science and Technology 38 (2004) 3933–3940. [10] A.L. Boreen, W.A. Arnold, K. McNeill, Triplet-sensitized photodegradation of sulfa drugs containing six-membered heterocyclic groups: identification of an SO2 extrusion photoproduct, Environmental Science and Technology 39 (2005) 3630–3638. [11] A. Albinet, C. Minero, D. Vione, Photochemical generation of reactive species upon irradiation of rainwater: negligible photoactivity of dissolved organic matter, Science of the Total Environment 408 (2010) 3367–3373. [12] J. Huang, S.A. Mabury, The role of carbonate radical in limiting the persistence of sulfur-containing chemicals in sunlit natural waters, Chemosphere 41 (2000) 1775–1782. [13] R.G. Zepp, P.F. Schlotzhauer, R.M. Sink, Photosensitized transformations involving electronic energy transfer in natural waters: role of humic substances, Environmental Science and Technology 19 (1985) 74–81. [14] A. Leifer, The Kinetics of Environmental Aquatic Photochemistry: Theory and Practice, American Chemical Society, Washington, DC, 1988. [15] Z. Manjun, Determination of photochemically-generated reactive oxygen species in natural water, Journal of Environmental Sciences 21 (2009) 303–306. [16] D. Vione, G. Falletti, V. Maurino, C. Minero, E. Pelizzetti, M. Malandrino, R. Ajassa, R.-I. Olariu, C. Arsene, Sources and sinks of hydroxyl radicals upon irradiation of natural water samples, Environmental Science and Technology 40 (2006) 3775–3781. [17] M.J. García-Galán, M.S. Díaz-Cruz, D. Barceló, Kinetic studies and characterization of photolytic products of sulfamethazine, sulfapyridine and their acetylated metabolites in water under simulated solar irradiation, Water Research 46 (2012) 711–722. [18] J.C. Anderson, J.C. Carlson, J.E. Low, J.K. Challis, C.S. Wong, C.W. Knapp, M.L. Hanson, Performance of a constructed wetland in Grand Marais, Manitoba, Canada: removal of nutrients, pharmaceuticals, and antibiotic resistance genes from municipal wastewater, Chemistry Central Journal 7 (2013) 54. [19] D. Dulin, T. Mill, Development and evaluation of sunlight actinometers, Environmental Science and Technology 16 (1982) 815–820. [20] P. Cardinal, Assessing nutrient and pharmaceutical removal efficiency from wastewater using shallow wetland treatment mesocosms. MSc Thesis. University of Manitoba, Winnipeg, 2013. [21] J.C. Carlson, J.C. Anderson, J.E. Low, P. Cardinal, S.D. MacKenzie, S.A. Beattie, J.K. Challis, R.J. Bennett, S.S. Meronek, R.P.A. Wilks, W.M. Buhay, C.S. Wong, M.L. Hanson, Presence and hazards of nutrients and emerging organic micropollutants from sewage lagoon discharges into Dead Horse Creek, Manitoba, Canada. Science of the Total Environment 445–446 (2013) 64–78. [22] R.F. Cross, J. Cao, Salt effects in capillary zone electrophoresis—III. Systematic and selective factors in the high ionic strength separation of sulphonamides in sodium phosphate buffers, Journal of Chromatography A 818 (1998) 217–229. [23] R.G. Zepp, Quantum yields for reaction of pollutants in dilute aqueous solution, Environmental Science and Technology 12 (1978) 327–329. [24] D.E. Latch, K. McNeill, Microheterogeneity of singlet oxygen distributions in irradiated humic acid solutions, Science 311 (2006) 1743–1747. [25] J. Huang, S.A. Mabury, Steady-state concentrations of carbonate radicals in field waters, Environmental Toxicology and Chemistry 19 (2000) 2181–2188. [26] S. Canonica, T. Kohn, M. Mac, F.J. Real, J. Wirz, U. Von Gunten, Photosensitizer method to determine rate constants for the reaction of carbonate radical with organic compounds, Environmental Science and Technology 39 (2005) 9182–9188. [27] D.E. Moore, W. Zhou, Photodegradation of sulfamethoxazole: a chemical system capable of monitoring seasonal changes in UVB intensity, Photochemistry and Photobiology 59 (1994) 497–502. [28] M.V.N. Mouamfon, W. Li, S. Lu, Z. Qiu, N. Chen, K. Lin, Photodegradation of sulphamethoxazole under UV-light irradiation at 254 nm, Environmental Technology 31 (2010) 489–494. [29] D. Fatta-Kassinos, M.I. Vasquez, K. Kümmerer, Transformation products of pharmaceuticals in surface waters and wastewater formed during photolysis and advanced oxidation processes—degradation, elucidation of byproducts and assessment of their biological potency, Chemosphere 85 (2011) 693–709.

J.K. Challis et al. / Journal of Photochemistry and Photobiology A: Chemistry 262 (2013) 14–21 [30] R.G. Zepp, J. Hoigne, H. Bader, Nitrate-induced photooxidation of trace organic-chemicals in water, Environmental Science and Technology 21 (1987) 443–450. [31] P.L. Miller, Y.-P. Chin, Indirect photolysis promoted by natural and engineered wetland water constituents: processes leading to alachlor degradation, Environmental Science and Technology 39 (2005) 4454–4462. [32] X. Zhou, K. Mopper, Determination of photochemically produced hydroxyl radicals in seawater and freshwater, Marine Chemistry 30 (1990) 71–88. [33] J.J. Guerard, P.L. Miller, T.D. Trouts, Y.-P. Chin, The role of fulvic acid composition in the photosensitized degradation of aquatic contaminants, Aquatic Sciences 71 (2009) 160–169. [34] M.W. Lam, S.A. Mabury, Photodegradation of the pharmaceuticals atorvastatin, carbamazepine, levofloxacin, and sulfamethoxazole in natural waters, Aquatic Sciences 67 (2005) 177–188. [35] A.G. Trovó, R.F.P. Nogueira, A. Agüera, A.R. Fernandez-Alba, C. Sirtori, S. Malato, Degradation of sulfamethoxazole in water by solar photo-Fenton. Chemical and toxicological evaluation, Water Research 43 (2009) 3922–3931.

21

[36] J. Wenk, S. Canonica, Phenolic antioxidants inhibit the triplet-induced transformation of anilines and sulfonamide antibiotics in aqueous solution, Environmental Science and Technology 46 (2012) 5455–5462. [37] J. Wenk, U. von Gunten, S. Canonica, Effect of dissolved organic matter on the transformation of contaminants induced by excited triplet states and the hydroxyl radical, Environmental Science and Technology 45 (2011) 1334–1340. [38] K. Kawabata, K. Sugihara, S. Sanoh, S. Kitamura, S. Ohta, Ultravioletphotoproduct of acetaminophen: structure determination and evaluation of ecotoxicological effect, Journal of Photochemistry and Photobiology A 249 (2012) 29–35. [39] R.S. Givens, B. Hrinczenko, J.H.-S. Liu, B. Matuszewski, J. Tholen-Collison, Photoextrusion of SO2 from arylmethyl sulfones: exploration of the mechanism by chemical trapping, chiral, and CIDNP probes, Journal of the American Chemical Society 106 (1984) 1779–1789. [40] A.G. Trovó, R.F.P. Nogueira, A. Agüera, C. Sirtori, A.R. Fernández-Alba, Photodegradation of sulfamethoxazole in various aqueous media: persistence, toxicity and photoproducts assessment, Chemosphere 77 (2009) 1292–1298.